Use of this information is subject to copyright laws and may require the permission of the owner of the information, as described in the ECHA Legal Notice.
EC number: 215-125-8 | CAS number: 1303-86-2
Relevance of terrestrial toxicity studies
The toxicity data on terrestrial organisms are from ecotoxicity tests that study relevant ecotoxicological parameters such as survival, growth, reproduction, and emergence. Relevant endpoints for soil microorganisms focused on functional parameters (such as respiration, nitrification, mineralization) and microbial growth. Enzymatic processes are considered not relevant for this risk assessment. The ecological relevance of enzymatic assays is questionable for several reasons:
§ The enzymatic activities are measured at conditions that are not representative for in situ conditions.
§ Several assays are conducted in pH buffered soil suspensions (some tests even at pH>10) and since the metal-enzyme interaction is pH dependent, this might obscure the relationship with effects in the soil.
§ Almost all assays use saturating substrate concentrations (typically several mM), a condition that is unlikely to occur in situ. The in situ effect of metals on an enzymatic reaction may be rather insensitive to the enzyme capacity (as measured with the enzyme assays) if substrate supply is the rate-limiting step.
§ The colorimetric reaction that is often required in enzymatic assays can also be subject to effects of metals (Nannipieri P. et al ,1997). and not all studies have experimentally verified this, i.e. some of the reported effects may be flawed
Relevancy of the test media
Only data from observations in natural and (OECD) artificial standard soil media have been used for the derivation of the PNEC.
The data used in the effect assessment should ideally be based on organisms and exposure conditions from Europe. This would, however, considerably reduce the amount of data that can be used. Therefore, data based on soils collected outside Europe have also been used. Two options may be followed here:
§ using all reliable data derived for non-European soil and
§ only using data for non-European soils with properties relevant for the EU conditions. These EU conditions are defined as 10thand 90thpercentiles of soil properties of the EU soils. These boundaries for the EU soils are based on the GEMAS-project (Geochemical Mapping of Agricultural and Grazing Land Soil project), which provides a harmonized and directly comparable dataset on soil properties (pH, organic matter content, clay content and effective CEC) and metals in 2211 samples of arable land soil (0-20 cm) and 2118 samples of grazing land soil (0-10 cm) at the EU scale (average sampling density of 1 site per 2500 km2, i.e. grid of 50 x 50 km). Only the ecotoxicity data were retained in case they were within the following boundaries: pH (0.01M CaCl2): 4.3 – 7.4; organic matter: 1.6 – 10.0% and clay: 6.0 – 37.0%.
Option 1 is selected because:
§ There are studies that show a tendency of increased boron toxicity in soils with low organic matter content, low clay content and pH < 7.5 (Aitken & McCallum, 1988; Gestring & Soltanpour, 1987; Van Laer et al., 2010). The effects found of soil properties on boron toxicity in soil are however rather limited (≤ factor 10) (see below),
§ Taking into account all soils, including the most sensitive ones is the most conservative approach
Relevancy of the test substance Studies on the ecotoxicity of boron have been performed with various compounds, such as boric acid (H3BO3), anhydrous sodium tetraborate (Na2B4O7), and hydrated sodium tetraborates (Na2B4O7.xH2O). For the purpose of this evaluation, all endpoints are converted to concentrations of elemental boron (B) using the relative molar mass.
Test duration What comprises “chronic exposure” is a function of the life cycle of the test organisms.A priorifixed exposure durations are therefore not relevant. The duration should be related to the typical life cycle and should ideally encompass the entire life cycle or, for longer-lived species the most sensitive life stage. Retained exposure durations should also be related to recommendations from standard ecotoxicity (e.g. ISO, OECD, ASTM) protocols. Typically chronic test durations for the higher plants are within the range of 4 (e.g. the root elongation test based on ISO 11269-1 (1995)) and 21 days (e.g. the shoot yield test based on ISO 11269-2 (1995)). OECD n° 208 (plant seedling emergence and growth test, 2006) recommended a test duration of between 14 and 21 days after emergence of the seedlings. Testing with soil invertebrates have a typical acute exposure duration of 7 to 14 days for the oligochaetesEisenia fetida/Eisenia andrei. Assessment of the chronic effects of substances on sub-lethal endpoints such as reproduction on oligochaetes has a typical exposure duration of 3 to 6 weeks for the standard organismEnchytraeus albidus(OECD, 2000; ISO 16387). For another standard speciesFolsomia candidasurvival and reproduction is typically assessed after 28 days of exposure (ISO 11267, 1999). Test durations using soil micro-organisms for the OECD 216 (carbon transformation test) and the OECD 217 (nitrogen transformation test) last 28 days..
Because boron is a necessary plant micronutrient, it is intentionally added in some instances where required by crop plants but is limited in the natural soil. Typical applied doses are 1-2 kg B/ha/yr, but may range from 0.5 to 7.6 mg B/ha depending on local conditions (Shorrocks, 1997; Borax 2002). This may be in the form of formulated fertilizers broadcast to agricultural soils or sprays applied directly to the plant or vicinity of the plants. In these instances, it might be appropriate to use a PNEC for agricultural soil that protects the agricultural uses of the soil, rather than a PNEC derived to protect non-agricultural or non-industrial soil. This is consistent with the REACH Guidance Document (2008) distinctions in developing PECs for agricultural, natural/grassland, and industrial soil. A potential approach would therefore be to derive a PNEC for agricultural soil based on toxicity, but also with consideration of the risk of deficiency. For natural soils, the presumption is that locally-adapted species will not be adversely affected by boron deficiency, so only boron toxicity is relevant for deriving a PNEC.
Boron is a naturally occurring element that is essential to a variety of organisms. In plants, it is necessary for a variety of metabolic processes (e.g. nitrogen metabolism, nucleic acid metabolism and membrane integrity and stability) and has been known to be an essential micronutrient for terrestrial plants for several decades (Butterwick et al., 1989; Eisler, 2000). Shorrocks (1997) documented the use of boron applications for 132 crops in over 80 countries, demonstrating the widespread nature of agricultural use of boron. Evidence exists that it is essential for nitrogen fixation in some species of algae (Smyth and Dugger, 1981), fungi and bacteria (Saiki et al., 1993; Fernandez et al., 1984), some diatoms and algae and macrophytes (Eisler, 2000). Required levels may vary, especially among plants, such that essential levels for one species may be toxic to another (Eisler, 1990).Work with rainbow trout and zebrafish has shown that embryo-larval development was adversely affected in waters deficient in boron (Rowe et al., 1998, Eckhert, 1998). Fort et al. (1998) reported that abnormal development in frog embryos (Xenopus laevis) was observed when larval stages were exposed to 0.003 mg-B/L or less. Boron does not appear to be essential for all species, however. Evaluation of essentiality in some animals is limited by the innate boron content in plant-based animal feeds. . The concentration-response curve for boron is likely to be U-shaped for most species, with adverse effects observed at high and low concentrations, while no adverse effects are observed at the intermediate concentrations (Lowengart, 2001).
Plant and animal species vary in the concentrations associated with deficiency and toxicity. Monocotyledons (e.g. corn and grasses) require about one-quarter as much boron as dicotyledons (e.g. tomatoes, carrots, clovers, beets) (Gupta 1985; Butterwick et al., 1989). The mobility of boron within the plant may help explain the observed deficiency and toxicity patterns. Boron is more mobile in plants that produce the simple sugars known as polyols (e.g. sorbitol and mannitol) than in species that do not produce polyols. In polyol-producing species, boron is translocated from one part of the plant to another and so may reach the meristem and affect growth. In the absence of polyols, boron is relatively immobile within the plant (Brown et al., 2002). A polyol-producing plant may be both more tolerant of boron deficiency and more sensitive to higher boron concentrations because of the mobility of boron within the plant. This is important in agricultural applications of boron, which may be applied as a soil treatment or as foliar spray. Agricultural application of boron depends on the plant and cultivar, as well as the local soil. Recommended application rates range from 0.5 to 7.6 kg B/ha (Borax, 2002), but typically are in the range of 1 to 2 kg B/ha (Shorrocks, 1997). If one assumes typical soil densities of 1700 kg/m³ and a mixing depth of 20 cm (default values used in the EUSES model), an application rate of 1 to 2 kg B/ha results in an estimated added soil concentration of 0.3 to 0.6 mg B/kg soil. Mortvedt et al. (1992) estimated soil concentrations of 0.16 to 2.0 mg B/kg soil for several crops with application rates of 0.45 to 5.7 kg/ha. The intentional application of borates to achieve such soil concentrations for agricultural crops should be acknowledged in the risk assessment process.
Chemistry of boron in soils
Boron may be considered a typical metalloid having properties intermediate between the metals and the electronegative non-metals. Boron has a tendency to form anionic rather than cationic complexes (Keren and Bingham, 1985). Boron does not undergo oxidation reduction reactions or volatilisation reactions in soils (Goldberg, 1997). Boron chemistry is of covalent boron compounds and not of B3+ions because of its very high ionisation potentials. Boron oxide, B2O3reacts with water to form boric acid, H3BO3. Boric acid is moderately soluble (4.9g 100mL-1water at 20°C). It acts as a weak Lewis acid by accepting a hydroxyl ion to form the borate anion. Aqueous boron species other than B(OH)3and B(OH)4-can be ignored for most practical purposes in soils (Keren and Bingham, 1985). In most soils with soil solutions in the pH range 4.0 to 9.0, the uncharged B(OH)3predominates. The borate ion is expected to form a variety of complex salts with suitable metal acceptor ions. However, there is relatively little evidence for the existence of metal borate complexes in solutions. Among the organic borates, the tendency is for boron to replace carbon or nitrogen in three-fold coordination (Keren and Bingham, 1985). In regions of low rainfall, the boron content of the soil is usually high. Boron in these soils probably exists largely as sodium-calcium borates. However, there is no information on the kinetics of dissolution of these minerals in water or on the composition of their products (Keren and Bingham, 1985).
Factors affecting the bioavailability of boron in soils
Boron toxicity to plants and many soil micro-organisms is a function of the bioavailability of the dissolved boron species in the soil solution and the ability of the soil to buffer boron concentrations in the soil solution. The bioavailability of metals and other inorganic substances in soil can be strongly affected by soil properties and slow equilibration reactions (ageing) after application to the soil. Various environmental factors can influence boron availability in soils, including pH, soil texture, organic matter content, soil moisture, and temperature. As boron is either neutral or negatively charged under environmentally relevant conditions, cation exchange capacity is not expected to play a relevant role. Investigations into properties, including ageing, which modify boron availability to plants are underway, but reports were not available in time to incorporate into this CSR. Boron availability to invertebrates depends on the relative amounts taken up by the organism by dermal adsorption and/or ingestion, although the relative importance of each route has not been determined (Vijver et al., 2001).
The amount of boron adsorbed by soil varies greatly with the contents of various soil constituents. Boron is adsorbed onto soil particles, with the degree of adsorption depending on the type of soil minerals present, pH, salinity, organic matter content, iron and aluminium oxide oxy/hydroxy content, and clay content (Hingston, 1964; Sims and Bingham, 1968; Bingham et al., 1970; Bingham, 1973). Boron adsorption can vary from being fully reversible to irreversible, depending on the soil type and environmental conditions (IPCS, 1998). As the pH is increased to about 9, the B(OH)4-concentration increases rapidly and the amount of adsorbed boron increases rapidly (Keren and Bingham, 1985). Hence, the critical range of extractable boron levels leading to deficiency is generally higher in alkaline soils than in acid soils (Bell, 1999). Boron reacts more strongly with clay than sandy soils (Keren and Bingham, 1985). Clay soils buffer boron in the soil solution better than sandy soils. The magnitude of the boron adsorption onto clay minerals is affected by the exchangeable cation (Keren and Gast, 1981; Keren and Mezuman, 1981; Keren and O'Connor, 1982; Mattigod et al., 1985). Calcium-rich clays adsorb more boron than sodium and potassium clays (Keren and Gast, 1981; Keren and O'Connor, 1982; Mattigod et al., 1985). Higher organic matter content increases the B-sorption capacity of soils (Yermiyahu et al., 1995). Adsorbed boron and boron adsorption maxima have been highly significantly correlated with organic carbon content (Elrashidi et al., 1982; Gupta, 1968). The uptake of boron by plants can be markedly affected by the presence of other plant nutrients in soils. The most well known of these is the effect of Ca (Gupta et al., 1985). There are few studies that compare boron toxicity for the same endpoint in different soils (Aitken & McCallum, 1988; Gestring & Soltanpour, 1987; Liang & Tabatabai, 1977; Liang & Tabatabai, 1978). The available results indicate a significant variation in boron toxicity thresholds among soils and show a tendency of increased boron toxicity in soils with low organic matter content, low clay content and pH < 7.5. The information is, however, too limited to allow conclusions on soil properties controlling boron toxicity in soils. Recently, Van Laer et al. (2010) studied the effect of soil type and ageing on the toxicity of boric acid to root elongation of barley seedlings in a set of 17 soils covering a large range in pH (4.4-7.8), organic carbon (0.14-30.7 %), clay content (2-59 %) and background boron concentration (1.2-32 mg B/kg). The EC10 values based on added boron concentration ranged between 3 and 27 mg B/kg dw among these soils. The percentage clay and organic carbon (log) were positively correlated with the log ED50 values of the root elongation test. Van Laer et al. also found positive correlations between %OC and %clay and % moisture content (%MC) of the soils. They suggest that increased moisture content of the soil dilutes added boron because adsorption is almost absent. Indeed, sorption of boron is very low: the solid-liquid distribution of added boron (Kd) ranges from 0 to 1.9 l/kg in the freshly spiked soils and hardly increases upon ageing. Further, %MC correlates well with ED50 (R2=0.67 on log scale and 0.76 on untransformed scale).
Ageing reactions The rate of boron adsorption on clay minerals is likely to consist of a continuum of fast adsorption reactions and slow fixation reactions. Short-term experiments have shown that boron adsorption reaches an apparent equilibrium in less than one day (Hingston, 1964; Keren et al., 1981). Long-term experiments have shown that fixation of boron increased even after six months of reaction time (Jasmund and Lindner, 1973). Studies on the residual effect of boron application after a single application also indicated decreasing boron toxicity to plants with increasing time since application (Gupta & Cutcliffe, 1984; Gestring & Soltanpour, 1987). Van Laer et al. (2010) studied the effect of ageing on the toxicity of boric acid to barley root elongation and microbial nitrification by measuring toxicity immediately after spiking soils and after ageing the soil for 1 or 5 months in closed containers (no leaching). Ageing at 20°C for up to 5 months in a closed container had only a slight effect on the boron toxicity: EC10 values increased by an average factor of 1.3 (range 0.4 – 3.5) after 1 month of ageing and 1.4 (range 0.5 – 3.3) in the 5 months. Because of this negligible effect of ageing on boron toxicity, it was decided to take into account the data for all equilibration times as replicates and calculate a geomean value for each soil for PNEC derivation. In contrast to this negligible ageing effect of added B, Van Laer et al. (2010) observed a large difference between soil partitioning between added boron and naturally present boron in these soils. Measured background soil solution boron concentrations in 3 soils were 0.2 to 0.9 mg B/l, equivalent to Kd values of 3 to 15 L/kg (Van Laer et al., 2010). Measured boron concentrations exhibiting effects on barley (EC10) in the added-boron soils ranged between 5 and 104 mg B/L across 17 natural soils (mean: 26 mg B/L). The median Kd value for added boron was 0.4 L/kg. Therefore, it was concluded that ageing after spiking for 5 months does not affect boron availability in the same way as natural geogenic or field equilibrated boron. The difference in boron speciation between added boron and native boron may be related to boron incorporation into silicate structures or boron in biomass. Van Laer et al. (2010) noted that the amount of boron naturally present in soils, as measured by aqua regia soluble boron, did not correspond to EC10 values based on added boron, even with ageing. They found the aqua regia-soluble boron to range from 1 to 32 mg B/kg in several natural soils, but these did not cause barley toxicity, even though EC10 values from freshly spiked soils ranged from 3 to 38 mg B/kg. This supports the used of added-boron, rather than total boron, as the basis for derivation of a PNEC for soil. Because of the large difference in bioavailability between boron naturally present in soils and added soluble B, risks of added soluble boron will be assessed by using the added risk concept.
Taking all information into account, it was decided not to implement normalization models for soil properties in the PNEC derivation for boron in soils because:
Information on Registered Substances comes from registration dossiers which have been assigned a registration number. The assignment of a registration number does however not guarantee that the information in the dossier is correct or that the dossier is compliant with Regulation (EC) No 1907/2006 (the REACH Regulation). This information has not been reviewed or verified by the Agency or any other authority. The content is subject to change without prior notice.Reproduction or further distribution of this information may be subject to copyright protection. Use of the information without obtaining the permission from the owner(s) of the respective information might violate the rights of the owner.
Welcome to the ECHA website. This site is not fully supported in Internet Explorer 7 (and earlier versions). Please upgrade your Internet Explorer to a newer version.
Close Do not show this message again